Henrique Miguel Veiga
Simão de Azevedo
Pereira
TESTES ECOTOXICOLÓGICOS COM CHIRONOMUS
RIPARIUS
ECOTOXICOLOGICAL TESTS USING CHIRONOMUS
Henrique Miguel Veiga
Simão de Azevedo
Pereira
TESTES ECOTOXICOLÓGICOS COM CHIRONOMUS
RIPARIUS
ECOTOXICOLOGICAL TESTS USING CHIRONOMUS
RIPARIUS
Dissertação apresentada à Universidade de Aveiro para cumprimento dos requisitos necessários à obtenção do grau de Doutor em Biologia, realizada sob a orientação científica do Professor Doutor Amadeu M.V.M Soares, Professor Catedrático do Departamento de Biologia da Universidade de Aveiro.
Apoio financeiro da Fundação para a Ciência e Tecnologia e do Fundo Social Europeu no âmbito do III Quadro Comunitário de Apoio através de uma Bolsa de Doutoramento com a
o júri
presidente Professor Doutor Joaquim José Borges Gouveia professor catedrático da Universidade de Aveiro
Professor Doutor Amadeu Mortágua Velho da Maia Soares (Orientador) professor catedrático da Universidade de Aveiro
Professor Doutor Rui Godinho Lobo Girão Ribeiro
professor associado com agregação da Faculdade de Ciências e Tecnologia da Universidade de Coimbra
Professor Doutor António José Arsénia Nogueira professor associado com agregação da Universidade de Aveiro Professor Doutor Fernando Manuel Raposo Morgado professor auxiliar com agregação da Universidade de Aveiro Professor Doutor José Vitor de Sousa Vingada professor auxiliar da Universidade do Minho
Doutor Carlos Barata Martí
investigador principal do Institute of Environmental Assessment and Water Research Jordi Girona – Barcelona, Espanha
Doutora Susana Patrícia Mendes Loureiro
investigadora auxiliar do Centro de Estudos do Ambiente e do Mar da Universidade de Aveiro
agradecimentos Este trabalho não teria sido possível sem a contribuição de diversas pessoas e entidades, a quem gostaria de apresentar os meus agradecimentos.
À Universidade de Aveiro, ao Departamento de Biologia e ao CESAM (Centro de Estudos do Ambiente e do Mar), pelas condições e meios proporcionados à realização dos trabalhos que conduziram à execução desta dissertação. À FCT (Fundação para a Ciência e Tecnologia) pelo financiamento prestado aos meus estudos, sob a forma de uma bolsa de doutoramento (SFRH/BD/18516/2004) e apoios associados.
Ao meu orientador, Professor Doutor Amadeu Soares, por me ter dado a oportunidade de realizar este trabalho, por toda a disponibilidade, orientação, apoio e paciência. Agradeço ainda a confiança que depositou em mim e a autonomia que me proporcionou ao longo destes anos, que em muito contribuiu para a minha aprendizagem.
Ao Professor Doutor António Nogueira por toda a ajuda, principalmente em relação à estatística; à Mónica e à Susana por estarem sempre disponíveis para ajudar.
Ao Professor Doutor José Paulo Sousa, pela ajuda na compreensão e utilização dos modelos cinéticos.
Ao Professor Doutor Jussi Kukkonen, por me ter recebido na Finlândia, pela amabilidade e partilha de experiências.
À Professora Doutora Lúcia Guilhermino e ao Doutor Carlos Gravato pela disponibilidade em me receber no seu laboratório e pela partilha de ideias. Ao pessoal do grupo que, ao longo destes anos, tornou a minha estadia no laboratório muito mais agradável, soube ter paciência para a minha ―excessiva‖ boa disposição e, muito particularmente, aos que tornaram o salão Riquinho uma referência no panorama da livre discussão de ideias entre amigos: Carla, Clara, Janeco, Fabi, Mariaki, Mimi, Pest(an)inha, Quim, Raquel, Sara, Serra, Vanessa, Zé. Um abraço especial ao Abel, por estar sempre disponível, ao Siz por toda a ajuda, e à minha companheira de viagem doutoral, Salomé. Ao Marco quero agradecer a constante motivação, ajuda e, acima de tudo, a amizade (do peito).
Aos amigos do Departamento, em especial ao Carlos Fonseca, à Catarina, à Cris, à Lísia e ao Nélson, pela simpatia e atenção que sempre me dedicaram. A todos os amigos que, embora não me acompanhassem directamente no trabalho e, mesmo que não o soubessem, sempre me levantaram a moral: Alex, André, Beto, Bez, Célia, Cê, Daniel, Hugo, Inês, Judy, Marco, Mequinha, Susana (força miúda, vais ultrapassar isso!), Teresinha e Tóia. Agradeço ainda aos Antigos Orfeonistas da UC os momentos de descontracção, boa disposição e amizade.
A toda a minha família, por aguentar os meus devaneios, e que tanto me ajudou mesmo quando não sabia que o estava a fazer. À minha avó Adélia, pela dedicação; à minha maninha Susana, pelo carinho.
À Elsa, por estar sempre comigo. O amor e apoio incondicional que me dedicas fazem-me ultrapassar todos os obstáculos. Nunca é demais referir: a importância que tens na minha vida é um reflexo do amor que sinto por ti. À minha mãe, pela força da natureza que é, pelos valores que sempre me transmitiu, pelo constante incentivo e, principalmente, pelo exemplo que
palavras-chave Mercúrio; Imidacloprid; Chironomus riparius; Toxicidade no desenvolvimento; Biomarcadores; Resposta comportamental
resumo Os sistemas aquáticos naturais podem estar sujeitos frequentemente a entrada
de tóxicos, quer seja através da lixiviação dos campos agrícolas ou da descarga por parte de unidades industriais. Avaliar o impacto potencial destes contaminantes nos sistemas aquáticos é muito importante, porque pode promover consequências sérias no balanço ecológico dos ecossistemas. Os efeitos de níveis sub-letais destes tóxicos nas populações aquáticas são detectados, em muitos casos, somente após diversas gerações, dependendo da espécie e do contaminante. O comportamento animal é considerado como sendo a primeira linha de defesa perante estímulos ambientais, e pode ser uma representação de alterações fisiológicas no organismo, sendo portanto um indicador excelente de alterações ambientais. O desenvolvimento dos sistemas de aviso prévio que integram parâmetros comportamentais pode ajudar a prever mais rapidamente possíveis alterações ao nível das populações naturais, do que a utilização de testes ecotoxicológicos padrão com a mesma finalidade. O conhecimento acerca de possíveis implicações devido a alterações comportamentais, em organismos bentónicos e em populações do campo sujeitas a tóxicos, é ainda escasso. Sabendo isto, neste estudo pretendeu-se investigar como o comportamento de Chironomus riparius – usando um biomonitor em tempo real – e outros parâmetros tais como crescimento, emergência de adultos, bioacumulação e biomarcadores, são afectados pela exposição a imidacloprid e ao mercúrio, que foram seleccionados como contaminantes. Os resultados demonstraram que a exposição às concentrações sub-letais de imidacloprid afecta o crescimento e o comportamento dos quironomídeos e que estes organismos podem recuperar de uma exposição curta ao insecticida. O comportamento que corresponde à ventilação de C. riparius revelou-se como um parâmetro mais sensível do que a locomoção e do que as respostas bioquímicas, quando as larvas foram sujeitas ao imidacloprid. Larvas de C. riparius expostas a concentrações sub-letais de mercúrio apresentaram uma tendência de diminuição de actividade comportamental, em testes com concentrações crescentes do tóxico; o crescimento das larvas foi também prejudicado, e as taxas de emergência de adultos e o tempo de desenvolvimento apresentaram retardamento. Estes organismos podem bioacumular rapidamente o mercúrio em condições de não alimentação e apresentam uma lenta depuração deste metal. Estes efeitos podem, em último caso, conduzir a prováveis repercussões ao nível da população e das comunidades. As reduções em actividades comportamentais, mesmo em concentrações baixas, podem diminuir a quantidade de tempo gasta na procura de alimento, produzindo efeitos aos níveis morfo-fisiológicos, e assim afectar severamente o desempenho dos quironomídeos no ambiente. O uso destes factores comportamentais como um parâmetro ecotoxicológico sub-letal relevante ao nível da toxicologia aumentará a versatilidade dos testes, permitindo uma resposta comportamental mensurável e quantitativa ao nível do organismo, utilizando uma avaliação não destrutiva, e assim certificando que esta aproximação pode ser usada em testes ecotoxicológicos futuros.
keywords Mercury; Imidacloprid; Chironomus riparius; Developmental toxicity; Biomarker; Behavioural responses
abstract Natural aquatic systems can be frequently subjected to toxicant inputs, either by runoff from agriculture fields or discharge from industrial plants. Assessing the potential impact of these contaminants on aquatic systems is an asset, as it can elicit serious consequences to the ecosystem balance. The effects of sub-lethal levels of these contaminants at the aquatic population levels are only detected, in many cases, after several generations, depending on the species and contaminant. Behaviour is considered to be the first line of defence towards environmental stimuli, and being a representation of physiological alterations in the organism it can be an excellent indicator of environmental changes. The development of early warning systems comprising behavioural endpoints can help to predict possible alterations at the field population levels even faster than conventional standard ecotoxicological tests.
The knowledge on the implication of behavioural disturbance by toxicants in benthonic organisms and in field populations is still scarce. Bearing this in mind, this study aimed to investigate how Chironomus riparius’ behaviour – using an online biomonitor – and other parameters such as growth, emergence, bioaccumulation and biomarker effect, are affected by exposure to the selected toxicants imidacloprid and mercury.
Results have shown that exposure to sub-lethal concentrations of imidacloprid affects growth and behaviour of chironomids and the organisms can recover from a short exposure to the insecticide. C. riparius ventilation behaviour appeared as a more sensitive endpoint than locomotion and biochemical responses when larvae were subjected to imidacloprid. Sub-lethal concentrations of mercury on C. riparius elicited a trend of impairment in behavioural patterns with increasing concentrations of the toxicant; growth was also impaired and delayed emergence rates / development time were noticed. These organisms can also quickly bioaccumulate mercury in unfed conditions and present a slow depuration of the heavy metal. These effects may in last instance lead to probable repercussions at the population and community level. Reductions in behavioural activities even at low concentrations might decrease the amount of time spent foraging, producing effects at the morpho-physiological levels, and thus severely affecting the chironomids performance in the environment.
The use of these behavioural endpoints as a sub-lethal ecotoxicogical relevant parameter in toxicology will increase the versatility of the tests, allowing a measurable and quantitative behavioural response at the whole-organism level in a non-destructive assessment, thus certifying that this approach can be used in further assays.
TABLE OF CONTENTS
TABLE OF CONTENTS ... v
LIST OF FIGURES ... viii
LIST OF TABLES ... x
1. GENERAL INTRODUCTION ... 1
1.1. Preamble ... 1
1.2. Behaviour as an endpoint ... 4
1.3. Selection of the test species ... 6
1.3.1. Chironomids: Ecology, biology and toxicology ... 6
1.4. Tested chemicals: Insecticide (Imidacloprid) and a heavy metal (Mercury) .. 9
1.4.1. Imidacloprid ... 9
1.4.2. Mercury ... 11
1.5. Research goals and thesis outline ... 14
References ... 16
2. BEHAVIOUR AND GROWTH OF CHIRONOMUS RIPARIUS MEIGEN (DIPTERA: CHIRONOMIDAE) UNDER IMIDACLOPRID PULSE AND CONSTANT EXPOSURE SCENARIOS ... 25
Abstract ... 26
2.1. Introduction ... 27
2.2. Material and Methods ... 29
2.2.1. Test organisms ... 29
2.2.2. Imidacloprid ... 30
2.2.3. Acute toxic experiments ... 31
2.2.4. Organisms’ exposure ... 31
2.3. Results ... 33
2.4. Discussion ... 37
References ... 41
3. EFFECTS OF IMIDACLOPRID EXPOSURE ON CHIRONOMUS RIPARIUS MEIGEN LARVAE: LINKING ACETYLCHOLINESTERASE ACTIVITY TO BEHAVIOUR ... 47
Abstract ... 48
3.1. Introduction ... 49
3.2. Material and Methods ... 51
3.2.1. Test organism ... 51 3.2.2. Imidacloprid ... 52 3.2.3. Organisms’ exposure ... 52 3.2.4. Behaviour ... 53 3.2.5. Biochemical analysis ... 53 3.2.6. Statistics ... 54 3.3. Results ... 54 3.4. Discussion ... 59 3.5. Conclusion ... 61 References ... 63
4. EFFECTS OF MERCURY ON GROWTH, EMERGENCE AND BEHAVIOUR OF CHIRONOMUS RIPARIUS MEIGEN (DIPTERA: CHIRONOMIDAE) ... 69
Abstract ... 70
4.1. Introduction ... 71
4.2. Material and Methods ... 73
4.2.1. Test organism ... 73
4.2.2. Test chemical ... 74
4.2.3. Water-only exposures: Range finding test / LC50 determination ... 74
4.2.4. Chronic experiments ... 75
4.2.5. Mercury analysis... 77
4.3. Results ... 78
4.4. Discussion ... 84
References ... 89
5. BIOACCUMULATION AND ELIMINATION OF MERCURY IN THE MIDGE LARVAE CHIRONOMUS RIPARIUS MEIGEN (DIPTERA: CHIRONOMIDAE): A LINK TO BEHAVIOUR. ... 97
Abstract ... 98
5.1. Introduction ... 99
5.2. Material and Methods ... 101
5.2.1. Test organism ... 101 5.2.2. Test chemical ... 102 5.2.3. Uptake experiment ... 102 5.2.4. Elimination experiment ... 104 5.2.5. Mercury analysis ... 105 5.2.6. Kinetics ... 105 5.2.7. Statistical analysis ... 106 5.3. Results ... 106 5.4. Discussion ... 108 References ... 112
6. GENERAL CONCLUSIONS AND FINAL REMARKS... 117
LIST OF FIGURES
Figure 1.1 - Relationship between behavioural ecotoxicology and other disciplines [adapted from Dell’Omo (2002)]. ... 5
Figure 1.2 - Chironomid life cycle, displaying the egg stage (1), larval stage (2), pupal stage (3) and terrestrial imago (4) [adapted from Ristola (2000)]. ... 7
Figure 1.3 - Structure of the synthetic insecticide imidacloprid [adapted from Matsuda et al., (2001)]. ... 9
Figure 2.1 - Average growth of Chironomus riparius when exposed to imidacloprid for a period of 96 and 240 h and when exposed for a period of 96 hours followed by a post-exposure period of 144 h in clean water. (a,b,c) same letters represent
differences between treatments at a significance level p<0.05 (ANOVA, Tukey’s
test). ... 34
Figure 2.2 - Average activity frequencies of locomotion of Chironomus riparius when exposed for a period of 4, 6, 8, and 10 days to imidacloprid (A), and when exposed for a period of four days to imidacloprid, followed by a post-exposure period of 2, 4, and 6 days in clean water (B). Concentrations used are expressed in the XX axis, as well as the days of recording in the Multispecies Freshwater Biomonitor. (*) represent significance level p<0.001 (ANOVA, Tukey’s test). ... 35
Figure 2.3 - Average activity frequencies of ventilation of Chironomus riparius when exposed for a period of 4, 6, 8, and 10 days to imidacloprid (A), and when exposed for a period of four days to imidacloprid, followed by a post-exposure period of 2, 4, and 6 days in clean water (B). Concentrations used are expressed in the XX axis, as well as the days of recording in the Multispecies Freshwater Biomonitor. (*) represent significance level p<0.001 (ANOVA, Tukey’s test). ... 36
Figure 3.1 - Average activity frequencies of locomotion of Chironomus riparius exposed to imidacloprid for a period of two and four days, followed by a post exposure period of two days in clean water. (*) represents significance level p<0.01 and (**) represents significance level p<0.001 in comparison with the control (ANOVA, Tukey’s test). (†) represents significance level p<0.01 for the comparison between times of exposure (ANOVA, Tukey’s test). ... 56
Figure 3.2 - Average activity frequencies of ventilation of Chironomus riparius exposed to imidacloprid for a period of two and four days, followed by a post exposure period of two days in clean water. (*) represents significance level p<0.01 and (**) represents significance level p<0.001 in comparison with the control (ANOVA, Tukey’s test). (†) represents significance level p<0.01 for the comparison between times of exposure (ANOVA, Tukey’s test). ... 57
Figure 3.3 - Average acetylcholinesterase activity, of Chironomus riparius exposed to imidacloprid for a period of two and four days, followed by a post exposure period of two days in clean water. (*) represents significance level p<0.01 and (**) represents significance level p<0.001 in comparison with the control (ANOVA, Tukey’s test). (†) represents significance level p<0.01 for the comparison between times of exposure (ANOVA, Tukey’s test). ... 58
Figure 4.1 - Mercury (ng Hg) fluctuation throughout the experimental period. A – ng Hg dynamic in water; B – ng Hg dynamic in sediment; C – ng Hg dynamic in biota. ... 80
Figure 4.2 - C. riparius growth measurements represented by body length at day 8 subtracted by the initial body length (mean + SD) after exposure to mercury chloride. Asterisks highlight treatments that are significantly different from the control (p<0.05). ... 81
Figure 4.4 - C. riparius mean emergence ratio (Mean ± SE). Asterisks highlight treatments that are significantly different from the control (p<0.05). ... 82
Figure 4.5 - Average activity frequencies of locomotion and ventilation of C.
riparius when exposed to mercury for a period of 4 days (A) and 10 days (B).
Vertical bars represent Standard Error (SE). Asterisks highlight treatments that are significantly different from the control (p<0.05). ... 83
Figure 5.1 - Kinetic behaviour of mercuric chloride in Chironomus riparius during uptake and elimination phases. Organisms were exposed to contaminated ASTM hard water during the first 24h. Data was fitted by nonlinear regression (see text for further explanation). ... 107
Figure 5.2 - Average activity frequencies of locomotion of Chironomus riparius
when exposed to a concentration of 31 µg L-1 [Hg] for a period of 72h. Dashed line
represents the end of the uptake phase and the beginning of the elimination phase. ... 107
Figure 5.3 - Average activity frequencies of ventilation of Chironomus riparius
when exposed to a concentration of 31 µg L-1 [Hg] for a period of 72h. Dashed line
represents the end of the uptake phase and the beginning of the elimination phase. ... 108
LIST OF TABLES
Table 4.1 - Grain size fractions of the inorganic acid-washed fine sediment. ... 75 Table 4.2 – ANOVA results... 79
Chapter 1
General Introduction
1. GENERAL INTRODUCTION
1.1. Preamble
Many surface water bodies are now contaminated due to the increasing usage of pesticides, mainly in agriculture, and to heavy metal contaminations from industry and/or natural sources. This contamination may cause impairment of ecological functions (Fleeger et al., 2003) and decline of non-target species (Rohr et al., 2006).
Ecotoxicology is interested in studying the effects of toxicants on the ecosystems. Pollutants matter because of their effects on populations and communities, through their effects on individual organisms (Moriarty, 1993). Since the immediate effects of pollutants are on organisms, either indirect (through habitat alterations) or direct (toxic effects of chemicals at the organismal level), one needs to assess what happens at the individual level to understand the impact on populations. The direct effect on individuals may range from rapid death through sub-lethal effects to no effects at all (Moriarty, 1993).
Ecotoxicology tests are needed to anticipate how toxicants are likely to impact ecological systems and to assess what changes are taking place in these systems under the influence of released toxic substances (Calow, 1997). When assessing the effects of a certain pollutant on a test species, the endpoints generally used are mortality (quantal type of data), and sub-lethal parameters like growth, reproduction, bioaccumulation and/or biomarker expression (continuous data), among others (Adams and Rowland, 2003). These responses can be ecologically relevant, as they are important components of fitness and determine the health, structure and dynamics of populations (Sibley et al., 1997).
In aquatic ecotoxicology the effects of anthropogenic (and natural) toxicants on aquatic biota are studied. These contaminants enter (and can be deposited) in the aquatic environment either from direct discharge from effluents, terrestrial runoff,
monitoring (identify unanticipated contamination and effects) or by targeted monitoring (focused on specific, known contaminant situations) (Grue et al., 2002). In the aquatic environments, biomonitoring may involve sampling of organisms as an indication of possible contamination, and in situ tests by assessing acute and chronic toxicity in caged organisms exposed to either contaminated water, sediment or both. Laboratory toxicity tests are also an asset, either by transposing and assessing in the laboratory field organisms and/or contaminated water/sediment; by using test species cultured in the laboratory with field contaminated water/sediment; or by assessing test species with artificially contaminated sediment/water. One must also bear in mind the ecological relevance of the experimental approach, in order to reach a compromise between realistic exposure situations and the scientific interest of the study.
Tested species can be representatives of the studied populations or model organisms that are regularly used in toxicity tests, with well studied endpoints. In this thesis the benthonic midge larvae of Chironomus riparius (Meigen) were used as model organisms. Several guidelines (e.g. EPA, 2000; OECD, 2004) are in use for this species, in order to standardize ecotoxicological tests and allow the replicability, repeatability and reproducibility of the experiments, thus increasing the test precision and uniformity among laboratories.
One of the drawbacks in using benthic macroinvertebrates for biomonitoring and assessment of water quality is the amount of effort required to process the samples, either in in situ tests (e.g. sorting animals in the sediment, measurements at the laboratory) or in laboratory tests where, in chronic tests, quantitative results on toxicity are only available at the end of the experimental period. For instance, to assess sub-lethal toxicity of pollutants on chironomids in laboratory, results on the effects on growth are only available after several days [e.g. 10 days and larvae still need to be measured (OECD, 2004)]; on emergence, after 20-28 days (OECD, 2004); on head-capsule deformity induction after several days [e.g. 10 days and still needs to mount the larvae (Meregalli et al., 2001)]; and to assess biomarker effects one needs to process the samples and quantify biomarker activities (Domingues et al., 2007), which can be time consuming. Survival is measured daily (one need to bear in mind that dead larva may be in the
sediment, thus not visible), but only at the end of the experimental period one can have certainties about mortalities. Ecotoxicologists want to use bioassays that are quick and easy, giving valuable information readily on contaminant effects on individual organisms, in order to make predictions about long-term impacts at an ecological level. In fact, Forbes et al. (2006) referred the need to devote more effort in developing and improving methods that directly measure effects of chemical impacts on populations, communities and ecosystems, and that less effort must be invested on measures that, at best, can only ever be suggestive of risks.
Beitinger and McCauley (1990) suggested that responses to environmental
changes can be divided in four categories: passive – no response, when the
stimulus is not sensed or occurs too rapidly thus leading to a decrease in performance capacities or even death; behavioural reactions – when subjected to certain chemicals, animals usually react in seconds or minutes, avoiding stress and trying to obtain a favourable position relative to the level of stimulus;
physiological responses – organisms suffer internal changes in various
physiological processes, including adjustments in physiological rate functions and tolerance acclimation enhancement, which may occur within hours to weeks; and
biochemical responses – synthesis of new molecules like ―stress‖ proteins in
response to environmental change, in order to restore homeostasis within genetic constrains, which may take from days to weeks. So, adding behaviour as an endpoint can help to formulate a quantitative minute-to-minute or hour-to-hour assessment of how tested species are (re)acting towards the toxicant concentration, bearing in mind that behaviour can be classified as the cumulative interaction of a variety of biotic and abiotic factors that represents the animal’s response to internal (physiological) and external (environmental, social) factors and that relates one organism to another (Dell’Omo, 2002). Behaviour provides an insight into various levels of biological organization, being a result and determinant of molecular, physiological, and ecological aspects of toxicology (Scott and Sloman, 2004). Therefore, behavioural responses may reflect biochemical changes in the individual organism and subsequently promote alterations in
communities, which can be translated into ecological consequences (Lagadic et al., 1994).
In former studies (e.g. with fish) behavioural parameters (considering swimming, ventilation, and foraging) have been suggested to be more sensitive than other endpoints (Beitinger, 1990; Beitinger and McCauley, 1990; Dell’Omo, 2002; Gerhardt, 2007). However, few studies have been made linking behavioural parameters to other biological (physiological, morphological) and ecological responses.
1.2. Behaviour as an endpoint
Behavioural responses comprise the first line of defence against adverse stimuli, since they can come into play within seconds after a stimulus is encountered (Beitinger, 1990). Behaviour can therefore be classified in different ways (Gerhardt, 2007): internal biochemical / physiological processes / mechanisms (neurobiological, hormonal, etc.); external ecological effects / consequences / purpose (e.g., avoidance, mating); degree of complexity such as from foraging behaviour; the distinction between individual (locomotion, foraging, learning with increasing complexity) and interactive behaviour (interspecific interactions such as predator-prey, or intraspecific interaction such as aggregation, territoriality, social interaction, reproduction related behaviours such as courtship, mating, spawning and parental care, etc.).
The study of behaviour in ecotoxicology, or behavioural ecotoxicology, is a comprehensive field which is the summation of many interactive disciplines, like ethology, ecology and toxicology (Fig. 1.1; Dell’Omo, 2002), studying how behaviour is modified by environmental pollutants.
The integrative nature of this parameter has some advantages: short response times (early warning responses), its non-invasive and non-destructive sensitiveness, and presents ecological relevance in laboratory toxicity tests (Depledge and Fossi, 1994; Gerhardt et al., 1994).
The objective of a behavioural bioassay is to determine whether a stimulus elicits an abnormal or adverse behavioural change outside the normal range of variability in an organism, which will adversely affect its survival, growth or reproduction (Beitinger, 1990). When an organism is subjected to an adverse stimulus, if there is no immediate physiological shock, usually that organism may behaviourally avoid the stimulus and effectively reduce exposure (Beitinger, 1990). If there is no avoidance capacity, organisms may suffer impairment of physiological responses, translated into a decrease in behavioural activities, making them less fit to avoid/hide from predators and/or foraging, thus increasing the lethality probability. The development of behavioural endpoints in toxicity assessment has improved the sensitivity and versatility of these studies, providing a unique toxicological perspective because they link biochemical causes of pollutants with ecological consequences on the population and community levels (Little, 1990).
Behaviour responses can be addressed by using avoidance tests, by empirical observation or by using biomonitors. Avoidance tests are mainly focused on ecological risk assessment using soil organisms (Loureiro et al., 2005; Natal-da-Luz et al., 2008). Empirical visual observation and manual data analysis is commonly used in toxicological tests to assess location of the animals during the
Ethology
Toxicology Ecology
Behavioural ecotoxicology
Figure 1.1 - Relationship between behavioural ecotoxicology
at the end of the experiment while visual observations, although applied during the tests, do not give us a measurable and discriminated response. Besides, these observations are very time-consuming and many times even impossible (e.g. benthonic animals that are inside the sediment may exhibit deleterious behaviour and cannot be observed). Behavioural biomonitors are employed to provide a visual and, therefore, measurable and quantitative behavioural response at the whole-organism level, offering an ecologically relevant, sensitive, fast and non-destructive assessment. There are several types of biomonitors that have been employed in multiple ecology and toxicology tests in the past decade. Most frequently used biomonitors in ecotoxicology experiments are: using video/image by computer-aided video tracking system [e.g. locomotor activity of isopods, in soil (Engenheiro et al., 2005)]; or using test chambers based on quadropole impedance technique [(Gerhardt et al., 1994) e.g. behavioural activities of benthic invertebrates (Gerhardt et al., 2005; Macedo-Sousa et al., 2007), including tests with chironomids (Janssens de Bisthoven et al., 2004)].
1.3. Selection of the test species
The selection of the studied species was based on the following criteria: easiness of handling and keeping the organisms in the laboratory under controlled conditions; the organisms must live in the sediment, but need to evidence drifting/swimming behaviour; evidence of early studies comprising measured behaviour in biomonitors; and need to be sensitive towards the selected toxicants.
1.3.1. Chironomids: Ecology, biology and toxicology
Chironomidae (Insecta, Diptera), frequently referred to as non-biting midges, are opportunistic tube-dwelling detritivores (Pinder, 1986) that play an important part in freshwater ecosystems, as they are ubiquitous and often dominate the benthic
communities of lotic and lentic environments, preferring eutrophic and organic enriched waters (Armitage et al., 1995; Vos, 2001).
They are able to invade habitats from where other species (e.g. competitors and predators) are often excluded (Pinder, 1986) and act many times as a major food source for other animals (Armitage et al., 1995; Rieradevall et al., 1995; Garcia-Berthou, 1999), playing an important part in bioturbation and nutrient cycling (Svensson and Leonardson, 1996) in these ecosystems.
Chironomids have a short life cycle (Fig. 1.2), comprising eggs, four larval stages, pupal stage (aquatic phases), and an adult stage (terrestrial/aerial phase). Adult females after swarming and mating lay the gelatinous egg batches (arranged helicoidally) on the water surface attached to several substrates, like plants or rocks. After the hatching, first instar larvae (white/transparent) are mainly pelagic. The following instars often inhabit the upper layer of the sediment, building protecting tubes from sediment particles.
The later instars are red coloured due to the presence of haemoglobin, that allows the midges to be tolerant to poorly oxygenated conditions (Ewer, 1942; Pinder, 1986) and increases the probability of a possible adaptation of the larvae to environmental changes, due to the response flexibility of haemoglobin (Ha and Choi, 2008). Larval stage can vary from few weeks to several years, being strongly related with temperature and food quality/quantity (Pinder, 1986). Pupal stage
1
2
3 4
Figure 1.2 - Chironomid life cycle, displaying
the egg stage (1), larval stage (2), pupal stage (3) and terrestrial imago (4) [adapted from Ristola (2000)].
Despite being benthonic organisms, larvae can be often found in the water column (Takagi et al., 2005) and can evidence drifting behaviour (Boothroyd, 1995), a probable strategy to colonize other areas, avoid predators or escape from contaminated sites. These midges exhibit characteristic motile activities like swimming, wholebody respiratory undulations and crawling. Both swimming and respiratory undulation are fast movements that involve body bending in a head-to-tail direction, while crawling combines the alternating use of the abdominal and prothoracic pseudopods as anchorage points, producing a form of locomotion analogous to caterpillar-looping (Brackenbury, 2000).
This study focused on Chironomus riparius that, along with C. tentans, are among some of the test species recognized as useful tools to study sediment toxicity (Ankley et al., 1994), being frequently used to assess the toxicity of natural (Péry et al., 2003; Faria et al., 2007) or spiked (Milani et al., 2003; Åkerblom et al., 2008) sediments. They can also be used in water only toxicity tests (Lydy et al., 2000; Stuijfzand et al., 2000) or in tests simulating contamination events either by aerial dispersion or runoff from agricultural fields (Agra and Soares, 2009; Pestana et al., 2009). For the assessment of pesticide toxic effects, several standardized methods using chironomids have been developed (ASTM, 2000; EPA, 2000; OECD, 2004). These multiple procedures and assessments are possible due to its multivoltine life cycle (Groenendijk et al., 1998); to the above mentioned widespread occurrence and ecological relevance; easiness to rear under laboratory conditions (Péry et al., 2002); and because during the larval development they are frequently in contact with the sediment (Goodyear and McNeill, 1999). Thus, these larvae have a high potential to play an important role as sentinel organisms in environmental monitoring (Choi et al., 1998).
The C. riparius Meigen larvae used in experiments were originated from our laboratory culture, which has been maintained for several years, genetically enriched episodically.
1.4. Tested chemicals: Insecticide (Imidacloprid) and a heavy metal (Mercury)
1.4.1. Imidacloprid
One of the most innovative and growing group of pesticides is the neonicotinoids (Tomizawa and Casida, 2003). These pesticides are chemically related to nicotine and epibatidine that are agonists of the natural nicotinic acetylcholine receptor (nAChR) (Matsuda et al. 2001). Imidacloprid (IMI) is a chloronicotinoid insecticide that belongs to this new class of pesticides and is already commonly and worldwide applied in order to control sucking insects in crops (Tomizawa and Casida, 2005; Tomlin, 2000).
Developed and patented by Bayer CropScience® AG (Monheim, Germany), IMI
[1-[(6-chloro-3-pyridinyl)methyl]-N-nitro-2-imidazolidinimine (C9H10ClN5O2)] (Fig. 1.3)
is very neurotoxic, acting via direct contact or ingestion and subsequent binding to the postsynaptic nAChRs of insects. IMI prevents acetylcholine from binding to the same receptor and, because it is not promptly degraded by acetylcholinesterase, promotes overstimulation of the insects’ nervous systems, causing tremors, lack of muscular coordination, decreased activity, desensitization and blocking of the receptors leading to modified behaviours and probable death of insects (Matsuda et al., 2001; Tomizawa and Casida, 2003). Neonicotinoids present higher selectivity factors for insects versus mammals, which is attributable to both target site specificity and detoxification. Nicotinoids (e.g. nicotine) are cationic and consequently selective for the mammalian nAChR, while neonicotinoids are not
Figure 1.3 - Structure of the synthetic insecticide
imidacloprid [adapted from Matsuda et al., (2001)].
Confidor®, Admire® and Gaucho® are some examples of commercial systemic insecticides that have IMI as the active ingredient and are used worldwide to control sucking insect pests, soil insects, termites, and some chewing insects, being also effective against adult and larval stages. In fact, IMI is used in urban areas to control turf pests in household lawns, parks, athletic fields, golf courses, etc., and this type of use appears to be increasing (CCME, 2007).
Nowadays the use of neonicotinoids is rising quicker than that of any other insecticides (e.g. organophosphates, pyrethroids) (Matsuda et al., 2001) and annual sales of neonicotinoids already account for 11%–15% of the total insecticide market (Tomizawa and Casida, 2005). This is mainly owed to their outstanding plant systemic activity and because the use of other neuroactive insecticides is declining due to selection of resistant insect strains and increasing restrictions based on human safety considerations (Tomizawa and Casida, 2005). Due to the boost in the use of IMI, it has been frequently detected in aquatic systems (surface and groundwater), especially during rainfall events and in shallow wells (CCME, 2007), increasing the awareness on possible effects of low concentrations in aquatic life. Since it’s applied in terrestrial habitats it can reach surface and ground waters via drift, leaching or dissolved runoff (Fossen, 2006; Gupta et al., 2002). IMIs’ physical-chemical characteristics might promote this contamination: persistent in soils [soil photolysis half life is 38.9 days and soil
aerobic half life is 997 days (ExToxNet, 1998)]; high solubility in water [0.514 g L-1
at 20 ºC and pH 7 (ExToxNet, 1998)]; has a long water and sediment half life [66 days (Sanchez-Bayo and Goka, 2005)], with a slow hydrolysis [half life >30 days, depending on formulation, pH and temperature (Sarkar et al., 1999), and can reach almost 1 year (CCME, 2007)], but with a fast aquatic photolysis [half life of
CONFIDOR® is 2.1h at λ=280 (Wamhoff and Schneider, 1999)]; low log Kow [low
octanol-water partition coefficient – 0.57 (ExToxNet, 1998), with a low
accumulation potential in aquatic species]. In natural field conditions,
concentrations ranging from 0.13 to 11.9 μg L-1 of IMI have been registered
(Phillips and Bode, 2004; CCME, 2007).
Nauen et al. (2002) refers that P450-monooxygenases are important in IMI detoxification and resistance development in insects, and studies with the aerial
insects Apis mellifera (Suchail et al., 2003) and Musca domestica (Nishiwaki et al., 2004) showed that IMI is metabolised very quickly and thoroughly, but its metabolites may extend the action of the insecticide.
Regarding toxicity, IMI can wield lethal and sublethal effects on non target species, being extremely toxic to aquatic invertebrates even at low concentrations. Daphnia
magna shows a 48 h LC50 varying from 17.36 mg L-1 (Song et al., 1997) until 85
mg L-1 (ExToxNet, 1998); for Lumbriculus variegatus, a 96h EC50 for
immobilization of 6.2 μg L-1 was reported by Alexander et al. (2007); for
chironomids, Stoughton et al. (2008) found a 96h LC50 for Chironomus tentans
using the formulated product of IMI (Admire®) of 5.40 μg L-1, whilst for Chironomus
riparius a 96h EC50 for mortality of 12.94 μg L-1 was reported by Pestana et al.
(2009).
The exposure to the active ingredient compared to commercial products might yield different levels of toxicity to aquatic organisms, generating diverse results as this could vary depending on the formulation of several products, endpoints and species tested (CCME 2007; Jemec et al., 2007; Stoughton et al., 2008).
1.4.2. Mercury
The Minamata and Niigata (Japan) incidents in the 1950s and 1960s focused worldwide attention and concerns on environmental mercury pollution, when many people were poisoned by methylmercury after eating fish and shellfish highly contaminated by mercury from direct industrial sources (Wiener et al., 2003; Ekino et al., 2007). Despite the actual imposed legislations have intended to minimize and eliminate mercury discharges to the environment in the last decades, mercury legacy in sediments and soils continues to be a worldwide concern.
Mercury (Hg) is a non-essential metal (Group B), showing lack of specific binding to organic ligands, and form strong covalent bonds (Mason and Jenkins, 1995), presenting a high toxicity to all biota. As an elementary substance, mercury is persistent and cannot be degraded into harmless products. It will therefore be
permanently recycled in the physical, chemical and biological processes in the environment (OSPAR, 2000).
This heavy metal occurs naturally and is very common in the environment, deriving from both natural processes and anthropogenic activities (Wolfe et al., 1998; EPA, 2001; Wiener et al., 2003). In case of natural activities, the major occurrence is derived from fallout of atmospheric gases from volcanic activity and geothermic emissions or emissions from deep-hydrothermal vents. Anthropogenic emissions of mercury, like mine tailings or industry, have since pre-industrial times resulted in a deposition rate increase by a factor of 2-10 in the most industrialized regions (Europe, North America, south-eastern China) during the last 200 years (Hylander, 2001). Due to its high volatility, it can also be dispersed via atmospheric transportation and deposited in other regions (Morel et al. 1998; Boening 2000) mainly as Hg(II) (EPA, 1997), being in this way available to biota even in regions far away from any pollution source. As so, two cycles are believed to be involved in the environmental transport and distribution of mercury: a global atmospheric circulation of elemental mercury vapour from sources on land to the oceans, while locally transport and distribution depend on the methylation processes of inorganic mercury from mainly anthropogenic sources (Boening, 2000).
Despite declining use, mercury has many applications, like extracting gold from ores (using liquid metallic mercury); treatment of diseases such as syphilis (until
the 20th century) and parasitic infections; as fungicides in agriculture (Clarkson et
al., 2003); use in chlor-alkali plants [manufacture of chlorine and caustic soda from brine for, amongst other applications, use in the food industry, textile production, cleaning agents, water treatment and pharmaceuticals, as well as intermediates in manufacturing other substances (OSPAR, 2009)], industry that produces nearly 90% of European anthropogenic Hg emissions to the atmosphere (Hylander, 2001).
The extensive past industrial use of the metal and its compounds together with widespread agricultural application of organochemicals has frequently resulted in serious contamination of surface water and sediments (Ullrich et al., 2001).
Mercury can occur in three valence states (0, +1 and +2) and may be present in various physical and chemical forms in the environment. For instance, can form
salts in two ionic states: mercury (I) and mercury (II). The environmental cycle of mercury has four strongly interconnected compartments (atmospheric, terrestrial, aquatic and biotic). The atmospheric compartment is dominated by gaseous
elemental mercury (Hg0); the terrestrial compartment is dominated by Hg (II),
sorbed to organic matter in soils. On the other hand, the aquatic compartment is dominated by Hg (II)-ligand pairs in water and Hg (II) in sediments, whilst the biotic compartment is dominated by methylmercury (including in the higher trophic levels of the aquatic food web). Mercury is highly reactive in the environment and cycles readily among these compartments (Wiener et al., 2003). Metallic mercury (Hg)
can be oxidized to mercury ions (Hg2+) which have a high affinity to sediments and
which are easily transformed in the environment into mercuric ions (OSPAR,
2000). Elemental mercury (Hg0) in surface waters occurs mainly from the reduction
of Hg (II) compounds by aquatic organisms, and oxidation of Hg0 can conversely
form Hg (II) (Ullrich, 2001). Since Hg0 is very volatile (Morel et al., 1998), it can be
readily lost from the aquatic environment at normal temperatures, playing an important part in the global Hg cycle, since its atmospheric transport in the vapor phase represents one of the major pathways of global deposition. The inorganic Hg forms, then again, can be transformed into methylmercury (MeHg) by chemical
processes and microorganisms like sulfate-reducing bacteria – methylation
(Wiener et al., 2003). Subsequent exposure of biota to the newly formed MeHg, a potent neurotoxin that is readily accumulated by aquatic biota due to its lipophilic and protein-binding properties, may pose a threat to humans and other fish-eating (Ullrich et al., 2001).
Contamination of biota from freshwater ecosystems by this heavy-metal is therefore a chronic and widespread environmental problem (Eisler, 1987; Boening, 2000). At high exposures Hg causes behavioural modifications, growth inhibition, reproductive impairment, decreased embryo/larval survival, and a variety of neurological and enzymatic dysfunctions in aquatic species (Zillioux et al., 1993). Mercury bioaccumulation usually starts by exposure through direct contact, breathing or by ingestion and subsequent retention on a tissue or organ (Fisher and Reinfelder, 1995) and can happen even at low concentrations, exhibiting high
1997). In fact, because it can be biomagnified through the trophic chain (Morel et al., 1998), mercury bioaccumulation can present an ecological risk. The majority of the toxicological studies assessing the effects of Hg in aquatic biota have, therefore, focused on bioaccumulation and trophic transfer of the heavy-metal (Mason et al., 1995; Wong et al., 1997; Vázquez-Núñez et al., 2007; Žižek et al., 2007).
Lethal and sub-lethal toxicological endpoints in freshwater organisms have also been reported. Boening (2000) describes several LC50 values: 96h LC50 for fish
ranges between 33 and 400 µg L-1; for Hydropsyche betteni, 96h LC50 is 2000 µg
L-1; for Daphnia magna, the 48h LC50 is 3610 µg L-1. Vidal and Horne (2003) also
reported a 96h LC50 of 0.18 mg L-1 for Tubifex tubifex, while Rossaro et al. (1985)
accounted for a 48h LC50 of 750 µg L-1 for C. riparius.
So, mercury can have deleterious effects on biota, is easily bioaccumulated and very persistent in the environment, with the nature and reactions of the Hg species determining the solubility, transport and toxicity of Hg in aquatic ecosystems.
1.5. Research goals and thesis outline
The information on the implications of the disturbance of complex behaviours by toxicants in benthonic organisms and in field populations is still scarce. The development and validation of behavioural tests to provide early warning information on these organisms’ behavioural reactions to the potential impact of contaminants (in this study: mercury and imidacloprid) is very important, especially during discharge or runoff periods, in order to evaluate potential effects on field communities.
This thesis aimed to investigate how Chironomus riparius’ behaviour (a new approach, using an online biomonitor) and other endpoints (e.g. growth, emergence, bioaccumulation and biomarker effect) are affected by exposure to the selected contaminants. Results are expected to improve the knowledge of the effects of imidacloprid and mercury exposure on benthic larvae, and to assess the sensitiveness of the endpoints chosen for this study.
In this research, the null hypothesis tested was that mercury and imidacloprid do not affect the normal behaviour nor compromise the development of Chironomus
riparius.
To address this, four separate chapters are organized in order to focus on these issues, followed by a general conclusion with final remarks.
Chapter two: ―Behaviour and growth of Chironomus riparius Meigen (Diptera:
Chironomidae) under Imidacloprid pulse and constant exposure scenarios‖, we focused on the effects of the insecticide on growth and behaviour of the chironomids when subjected to a constant and a pulse (followed by a recovery period) exposure to the pesticide.
Chapter three: we address the ―Effects of imidacloprid exposure on Chironomus
riparius Meigen larvae: linking acetylcholinesterase activity to behaviour‖. In this
study we perform a link between parameters with ecological relevance at individual level (behavioural parameters) with biochemical responses, to fully understand xenobiotics’ mode of action.
Chapter four: ―Effects of mercury on growth, emergence and behaviour of
Chironomus riparius Meigen (Diptera: Chironomidae)‖, we assessed the effects of
mercury on C. riparius, simulating a mercury discharge. Growth was measured after 8 days exposure, while behaviour was measured at days 4 and 10, and emergence and development time were also assessed.
Chapter five: ―Bioaccumulation and elimination of mercury in the midge larvae
Chironomus riparius Meigen (Diptera: Chironomidae): a link to behaviour‖,
mercury toxicokinetics (uptake and elimination) was evaluated using C. riparius under a water-only exposure. Behavioural parameters were monitored during the experimental (uptake and elimination) period.
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